ASTUDY CONDUCTED FOR THE
Draft release to internal group: June 2010
External review: Dec 2010
External release: Feb 2011 Study Authors: Elsa Olivetti Jeremy Gregory Randolph Kirchain Massachusetts Institute of Technology Materials Systems Lab Phone: (617) 253‐0877 Fax: (617) 258‐7471 E‐mail: firstname.lastname@example.org
Approximately 80% of portable batteries manufactured in the US are so‐called alkaline dry cells with a global annual production exceeding 10 billion units. Today, the majority of these batteries go to landfills at end‐of‐life. An increased focus on environmental issues related to battery disposal, along with recently implemented battery directives in Europe and Canada and waste classification legislation in California, has intensified discussions about end‐of‐life battery regulations globally. The logistics of battery collection are intensive given the large quantity retired annually, their broad dispersion, and the small size of each battery. Careful evaluation of the environmental impacts of battery recycling is critical to determining the conditions under which recycling should occur. This work compares a baseline scenario involving landfilling of alkaline batteries as municipal solid waste with several collection schemes for battery recycling through pyrometallurgical material recovery. Network models and life cycle assessment methods enable the evaluation of various end‐of‐life collection and treatment scenarios for alkaline batteries.
The study employs life‐cycle assessment techniques in accordance with the ISO 14040 standard. This approach is applied to each end‐of‐life scenario to provide a comprehensive means of accounting for environmental impacts. The scope of the analyses includes raw material extraction and refining, battery manufacturing, end‐of‐life disposition, and transportation. Secondary life cycle inventory datasets from the ecoinvent 2.2 data set were employed when primary data were not available. Metrics evaluated include Cumulative Energy Demand (CED), Global Warming Potential (GWP), and Ecosystem Quality, Human Health, and Resources indicators (the latter three are damage categories from the Ecoindicator 99 methodology). Because of the country‐specific grid mixes used in the recycling scenarios, the amount of radioactive waste is also presented as a metric of interest in for the recycling scenarios.
To summarize the full life cycle implications of alkaline batteries, the production of raw materials dominates the life cycle with the transport of those raw materials to manufacturing having a minimal environmental impact as shown in the figures below using the proxy environmental impact metric, CED. A few materials dominate this materials production impact, with manganese dioxide, zinc, and steel having the highest impacts.
LIFE CYCLE IMPACT USING CED FOR 1 KG SALES‐WEIGHTED AVERAGE ALKALINE BATTERIES INCLUDING PACKAGING (~30 BATTERIES) PRODUCTION IMPACT (LEFT GRAPH) AND MATERIALS IMPACT (RIGHT GRAPH) USING CED FOR 1 KG SALES‐ WEIGHTED AVERAGE ALKALINE BATTERIES INCLUDING PACKAGING (~30 BATTERIES) There are complex and uncertain potential impacts associated with placing primary alkaline batteries in landfills at end‐of‐life and recycling may reduce those impacts, but may cause additional burdens that outweigh benefits. The primary factors that drive the environmental impact of alkaline battery recycling, compared to the baseline landfill scenario, include the recycling technology used, the amount of materials recovered, and the state of the recovered materials. Study findings indicate recovering more than zinc for metal value (replacing virgin material) is important for reducing environmental impact and technologies involving high temperature are energy intensive. The principal drivers of end‐of‐life environmental performance of batteries vary depending on the metrics of impact assessment. Findings indicate metrics around energy and carbon are strongly dependent on recovery technologies, metrics for ecosystem quality depend on landfill scenario assumptions as well as the materials benefits associated with recycling. For the latter, if one assumes little to no landfill leachate resulting from batteries (in other words, batteries remain intact in the landfill or leachate is collected and not of concern over the time horizon considered), the main benefit from recycling stems from the recovery of zinc, manganese and steel. The same is true for metrics around human health. Several conclusions related to the transportation of the batteries to their end‐of‐life disposition step are also of interest. For the recycling scenario where batteries are dropped off by individual consumers at a retail or municipal
facility, the assumed allocation of the trip (dedicated versus non‐dedicated) drives the burden. In addition, the fact that there are just three facilities modeled for North America that are able to take alkaline batteries for recycling drives the large transportation burden associated with taking batteries from an intermediary consolidation facility to the recycling facility. This study modeled several scenarios for collection and recycling of batteries that resulted in an overall net environmental burden as compared to landfilling as well as several that resulted in an overall environmental benefit. When considering metrics related to energy or global warming potential, the recycling scenarios appear more environmentally burdensome whereas for metrics of human health and ecosystem quality, the recycling scenarios appear more environmentally beneficial. This study does not intend to explicitly compare the technologies; rather this work investigates the specific sites and contexts under which the technologies operate. For the collection burden associated with battery recycling, the greatest burden was associated with the scenario wherein individual consumers dropped batteries off at municipal locations, such as transfer stations. The crucial assumptions were around the degree of dedication for this leg of the journey, and literature indicated a higher likelihood of a dedicated trip for municipal drop‐off along with greater distances traveled than the retail drop‐off. Municipal drop‐off was on average 3‐4 MJ and 0.2 kg CO2 eq/kg of batteries disposed greater than retail drop‐off (1.3x10‐7 DALY, 0.015 pdf*m2yr, and 0.25 MJ surplus /kg surplus for the other metrics). Curbside pickup (both MSW and Recycle co‐collection) for the recycling scenarios was determined to be lower in impact than both of the drop‐off scenarios.
For the materials recovery burden at the end‐of‐life, the overall conclusions indicated that the burden or benefit of recycling depends on the scenario assumed in recycling. The materials recovery credit associated with zinc often drives the environmental benefit of all the recycling scenarios, so investigating how this material acts within the metals market would offer further insight to its benefit. Five overall scenarios were examined with their specific transportation, fuel mix and materials recovery contexts. Scenario A, involving a metal fuming furnace in the Northwest, employed coal and a hydropower electrical grid to recover zinc for metal value with steel and manganese dioxides reporting to slag which is sold to cement market. It exhibits an environmental benefit as compared to the baseline MSW landfilling scenario for metrics of ecosystem quality for both municipal and retail drop‐off. This scenario results in a more significant environmental burden than landfilling as measured by CED, GWP, human health, and resources using any modeled collection method.
Scenario B, located in the Midwest used natural gas and coal, to recover zinc and steel to metal value with some metal value from manganese dioxides and the remainder reporting to slag which is sold to cement market. This scenario exhibits an environmental benefit or neutrality as compared to the baseline MSW landfilling scenario for metrics of ecosystem quality and human health for both municipal and retail drop‐off. However, this scenario results in a more significant environmental burden than landfilling as measured by CED, GWP, and resources using any modeled collection method.
Scenario C recovered steel to metal value and zinc and manganese dioxides sold for micronutrient replacement using a lower temperature process powered by natural gas and a Canadian electrical grid. It exhibits an environmental neutrality or slight benefit as compared to the baseline MSW landfilling scenario for metrics of human health and ecosystem quality. For CED, ecosystem quality, and resources
the value of this scenario is generally environmentally burdensome. Therefore, for the scenarios currently used in the US, the impact oscillates between environmentally beneficial and environmentally burdensome depending on the indicator metric investigated and the transportation scenario used. There were two scenarios investigated that are not currently used in the US. The first of these scenarios is modeled after using the EAF infrastructure in the US where zinc and steel (with volume of manganese) are recovered for metal value and exhibit an environmental benefit compared to a baseline landfilling scenario for all metrics used in this study except for an apparent neutrality in the case of CED and in municipal drop‐off for ecosystem quality. Because of the copper poisoning to steel production, this scenario is limited by capacity (although based on the number of batteries recovered, this value would not be exceeded). EAF facilities would require permitting for this scenario to be possible for EOL battery processing in California. This scenario is sensitive to zinc recovery as shown in the scenario analysis. The sensitivity analyses indicate that for the metrics of CED and GWP the burden for recycling with less than 32‐40% zinc recovery exceeds the impact of landfilling.
The final scenario, not currently used in the US, recovers zinc, steel and manganese for metal value based on European recycling facilities incorporating low carbon intensity electrical grids from France and Switzerland. The transportation scenario assumes transport by road and ship to the EU. The results indicate for the majority of cases, there is an environmental benefit to this scenario, except for CED and GWP where it is environmentally burdensome, and for resources where the transportation scenario (retail versus municipal drop‐off) dictates whether it is a burden or benefit.
When all the examined recycling scenarios are assumed to use the US electrical fuel mix as an electricity source and an average, equivalent transportation burden, Scenario D may demonstrate environmental benefit of recycling compared to landfilling while Scenario A, C and E are more environmentally burdensome and the environmental benefit of Scenario B is metric dependent. For ecosystem quality all the scenarios are environmentally beneficial except for C.
Incineration, as part of MSW management, performs similarly to the hypothetical EAF scenario, except that it is burdensome due to reduced materials recovery and increased transportation burden. Incineration may be preferable to landfilling because of the potential for material recovery.
ONTENTSExecutive Summary ... 2 Chapter 1: Introduction ... 7 Chapter 2: Introduction to life cycle assessment... 9 Chapter 3: Overall Goal and scope definition ... 13 Section 1: Whole Life Cycle Assessment ... 15 Chapter 4: Alkaline battery life cycle assessment ... 15 Methodology ... 15 Scope ... 15 Data sources and assumptions ... 16 ecoinvent data gaps ... 20 Data Quality/Source Matrix ... 20 Results ... 21 Section 2: Alkaline Battery end of life focus ... 29 Chapter 5: End‐of‐Life investigations ... 29 Goal, Scope and Methodology ... 29 Spent Battery Chemistry ... 30 Logistics assumptions ... 31 Baseline and Curbside Pickup scenarios ... 31 Consumer Drop‐off scenarios (municipal/retail) ... 33 Previous work ‐ logistics ... 35 Landfilling and Incineration Toxicity issues ... 36 Incineration ... 37 Landfilling ... 40 Recycling technologies ... 42 Scenario A ... 43 Scenario B ... 44 Scenario C... 44 Scenario D modeled after recycling with steel in an EAF ... 44 Scenario E modeled after aggregating European recyclers ... 45
Chapter 6: End of life scenario analysis results... 48 Incineration scenario ... 63 Chapter 7: Parameter Analysis... 64 Conclusions ... 72 Recommendations for actions to reduce the environmental impact ... 76 Life cycle impact ... 76 End of life impact ... 76 Future Work ... 76 References ... 78 Appendix A: NEMA survey questionnaire ... 81 Appendix B: Detail around LCI ... 84 Appendix C: Numeric results of Chapter 6 ... 93 Appendix D: Figures for 100% Primary offset for zinc and steel. ... 95 Appendix E: External Reviews ... 102 Reviewer 1 ... 102 Reviewer 2 ... 104 Reviewer 3 ... 108
End‐of‐life issues for consumer non‐durables has become increasingly subject to the critical eye of individuals, local government and producers. Most products follow a linear lifecycle, beginning as raw materials in the earth, passing through refining, manufacturing, and use, and finally returning to the earth in a landfill. While this linear lifecycle has been the norm for many products in the US, increasing focus on environmental issues has drawn attention to the apparent wastefulness of a linear lifecycle. To address this, adding loops to the linear lifecycle, often in the form of reuse, remanufacturing, and recycling, has been proposed. However, these loops, and in particular the recycling loop, are not without debate, as the economic and environmental impacts of such loops are often uncertain.
In the case of alkaline batteries, the economic and environmental impacts of recycling are particularly interesting. In the US, the large quantity of alkaline batteries that are retired each year, the broad dispersion of those batteries, and the small size of each individual battery, makes the logistics of battery collection particularly challenging. The material composition of alkaline batteries adds another layer of complexity to the recycling dilemma, as there are disparate views about whether materials found in alkaline batteries are harmful when landfilled. Although alkaline batteries pass all U.S. EPA hazardous
waste criteria and are therefore not deemed to be hazardous in the U.S., the state of California has deemed them harmful, and some consumers are under the impression that alkaline batteries contain harmful materials1. Given these and other issues, careful evaluation of both the economic and environmental impacts of alkaline battery collection and recycling is critical prior to deciding whether or not alkaline battery recycling should take place and, if so, under what conditions. In the US, clearly understanding such impacts is particularly relevant, as an increased focus on environmental issues, along with a recently implemented end‐of‐life battery directive in Europe and regulatory interpretation in California impacting alkalines, has intensified the discussions about end‐of‐life battery regulations in the US.
As background to the legislation that impacts batteries in the US, the situation in California presents one perspective. Most batteries are considered hazardous waste in California when they are discarded including batteries of all sizes, both rechargeable and single use. Therefore alkaline batteries, as of February 8, 2006, must be recycled, or taken to a household hazardous waste disposal facility, a universal waste handler, or an authorized recycling facility. Large and small quantity handlers are required to ship their universal waste to another handler, a universal waste transfer station, a recycling facility, or a disposal facility. Several other states in the US have legislated a restriction on disposal in landfills of particular rechargeable chemistries such as nickel cadmium, but these do not cover alkaline single use batteries and as of this report writing no other state has legislation banning alkaline batteries from landfills. Canada and the EU mandate collection of alkaline batteries.
This study evaluates the environmental impacts of different end‐of‐life strategies, such as disposal and recycling for alkaline batteries in the United States. The analysis is divided into two sections. The first section of the analysis encompasses the entire life cycle of the battery, accounting for impacts from production in a manufacturing facility to use and eventual end‐of‐life treatment. The focus of the second section of the study is more detail on the end‐of‐life treatment. The scope and approach are outlined in more detail below and in the two sections of the document. The geographic scope of the study includes batteries manufactured and disposed of in the United States. The results of the analysis were generated in accordance with the ISO 14040 standard for life cycle assessments (LCAs).
For the second section of the study, defining multiple scenarios enabled investigation of the implications of several end‐of‐life treatments for battery recycling. These consist first of a baseline scenario including municipal solid waste (MSW) pickup of batteries with regular household waste accompanied by disposal in a “typical” landfill. This baseline was contrasted with a series of recycling scenarios that included multiple collection schemes and recycling technologies. The collection schemes included curbside with MSW, curbside with municipal recycling, and drop off to both municipal and retail locations. Incineration of batteries with regular household waste will also be commented on throughout this document; however it is not a major focus because of the dominance of landfilling in the US, as shown in Figure 1.
There is no mercury added to US OEM-produced alkaline batteries, but may be present in trace quantities from other sources such as batteries produced before mercury was not added, imported or counterfeit batteries.
FIGURE 1. MUNICIPAL SOLID WASTE MANAGEMENT FROM 1960 TO 2007. REPRODUCED FROM  The analysis aimed to quantify the total vehicle miles traveled (VMT) by the batteries as a function of collection scheme through a series of network models and to measure the burdens and benefits of treatment at end of life through life cycle assessment techniques. The results describe the impact and the sensitivity of the analysis to several components including the effect of collection scheme, the effect of regional variation in VMT, and the energy needed to treat at end‐of‐life. Through these scenario analyses, this work explores some of the condition (or conditions) under which recycling demonstrates environmental benefit compared to landfilling. This document begins with a brief introduction to the LCA methodology, followed by the goal and scope definition of the overall study. The first section describes the scope, methodologies and results of the full LCA for alkaline batteries, and then an investigation into the end‐of‐life alternatives is shown in the second section along with a discussion of the sensitivity of the results to several key assumptions. The document concludes with a summary of the study results and recommendations for actions that may be taken to reduce the environmental impact associated with the products.
NTRODUCTION TO LIFE CYCLE ASSESSMENT
Life cycle assessment is an approach to analyzing the environmental impact of a product or industrial system throughout its entire life cycle, from cradle to grave. The life cycle under consideration generally encompasses all stages of a product’s life, including raw material production, product manufacture, use, and end‐of‐life disposal or recovery, as depicted in Figure 2. The arrow in Figure 2 demonstrates the transport that takes place between each phase in the life cycle. The comprehensiveness of LCA is one of its strengths; it includes many details that are not part of more focused environmental impact analyses.
However, the complexity and level of detail necessitate a strict adherence to a consistent methodology. A brief overview of the LCA methodology is presented here; more thorough references are available for additional details [2, 3]. Materials Production Manufacture and Assembly Use Recovery / Recycling FIGURE 2. PHASES IN A PRODUCT LIFE CYCLE. The International Organization for Standardization has developed a standard methodology for life cycle assessment as part of its ISO 14000 environmental management series. The LCA standard, ISO 14040 , outlines four main steps in an LCA: goal and scope definition, inventory analysis, impact assessment, and interpretation of results. These steps are shown in Figure 3, and explained below.
Goal & Scope Definition Inventory Analysis Impact Analysis In te rpre ta tio n FIGURE 3. LIFE CYCLE ASSESSMENT METHODOLOGY (ADAPTED FROM THE ISO 14040 STANDARD). • Goal and scope definition articulates the objectives, functional unit under consideration, and
regional and temporal boundaries of the assessment.
• Inventory analysis entails the quantification of energy, water, and material resource requirements and emissions to air, land, and water for all unit processes within the life cycle, as depicted in Figure 4.
• Impact assessment evaluates the human and ecological effects of the resource consumption and emissions to the environment associated with the life cycle.
• Interpretation includes an evaluation of the impact assessment results within the context of the limitations, uncertainty, and assumptions in the inventory data and the scope. FIGURE 4. INVENTORY ANALYSIS: INFLOWS AND OUTFLOWS OF A UNIT PROCESS. The accumulation of the life cycle inventory step transforms a detailed list of all the inputs and outputs of the process to and from the technosphere into inputs from and outputs to nature, therefore containing only resources and emissions.. The impact assessment step is particularly challenging because of the difficulty associated with aggregating and valuing numerous types of resource consumption and emissions to the environment. There is uncertainty in the modeling used to produce midpoint and endpoint indicators. There are metrics that include a single attribute such as energy or global warming potential, but there are also metrics that try to capture multiple impacts. The impact assessment methods (and proxy metrics) that were used in this study are described below.2
Single Issue metrics:
• Cumulative Energy Demand (CED) includes all direct and indirect energy consumption associated with a defined set of unit processes. It does not directly account for the impact of non‐energetic raw material consumption or emissions to the environment. Values for CED are measured in terms of energy (e.g., joules). Note: CED is a proxy metric and not a formal impact assessment method. This method considers energy from multiple sources, including renewable and non‐renewable. For the CED results, all energy sources are presented as both renewable and non‐renewable are used. This number is broken down by source where it is of interest. 
• Global Warming Potential (GWP) incorporates the impact of gaseous emissions according to their potential to contribute to global warming based on values published in 2007 by the Intergovernmental Panel on Climate Change. The impacts for all gaseous emissions are evaluated relative to carbon dioxide. Impact assessment values for GWP are measured in terms of an equivalent mass of carbon dioxide (e.g., kg CO2 equivalent) .
Multi impact metric:
For the recycling technologies, the amount of nuclear waste generated is also presented as a result because of the grid mixes used for some of the technologies.
• Ecoindicator 99 (EI) is a damage‐oriented method that calculates environmental impact in three categories: damage to human health, ecosystem quality, and damage to resources. The characterization of damage by inventory items is based on scientific methods (e.g., effects of toxic materials in drinking water on human health) and these damage categories serve as endpoints in ISO 14040. Damage to human health is in units of disability adjusted life years (DALY), implying that different disability caused by diseases are weighted. Ecosystem quality is reported in units of potentially disappeared fraction of plant species (PDF*m2yr). The final category is Resources, which includes assessment of minerals and fossil fuels in units of MJ surplus, or the additional energy requirement to compensate for future ore grade. Damage is then normalized by average European impacts. The egalitarian perspective was used for this study. The valuation of the relative importance of various environmental impact categories is determined by survey responses from an expert panel; these responses determine the relative weighting of all of the impacts. These steps enable all of the impacts to be aggregated into a single value which has the units of “points”, where 1000 points represents the average environmental impact of a European in one year. This European weighting may present a limitation given the US geographic scope of this analysis; however, this impact assessment method is used primarily for its damage categories rather than the single combined metric for this study. Also, because this study examines various options for end of life and is therefore classified as comparative assertion, weighting results are not used for communication. A few different approaches exist, however for the purposes of this study, the damage category impacts are reported including human health,
ecosystem quality and resources .
These methods were used for this work because they provide separate lenses with which to evaluate environmental burden. Global warming potential is of interest to the audience because of the focus on greenhouse gases in some pending legislation. Because of the potential ecosystem quality and human health concerns associated with batteries in landfills, the EI 99 method was also chosen. It can be challenging to have a “feel” for reasonable values calculated by life cycle impact assessment methodologies as such values often represent abstract concepts or non‐physical quantities. Indeed, such values are most useful when presented in the context of a comparison so that relative quantities may be evaluated. Several products and processes have been evaluated using the methodologies for LCAs detailed above, and the results appear in Table 1. These values provide some basis of comparison for the results presented in this study.
TABLE 1. LIFE CYCLE IMPACT ASSESSMENT VALUES FROM FOUR LCAS. VALUES CALCULATED USING THE ECOINVENT 2.2 DATABASE.
Product or Process CED
(MJ) GWP (kg CO2 eq) Human Health (DALY) Ecosystem quality (PDF*m2yr) Resources (MJ surplus) Production of 25 g PET beverage bottle (20 fl oz/590 ml) 2 0.07 6.8 x 10‐8 0.005 0.15 Production of 14 g aluminum beverage can (12 fl oz/350 ml) 3 0.2 2.6 x 10‐7 0.007 0.17 100 km fuel consumption in a European passenger car 305 18 2 x 10 ‐5 1 21 Coffee pot: 5 years 5400 220 2 x 10‐4 30 190 Now that the concept of life cycle assessment has been briefly outlined, the next section will define the goal and scope of the work undertaken in this study.
OAL AND SCOPE DEFINITION
This life cycle assessment is divided into two sections. The goal of the first section, described in Chapter 4, was to determine the life cycle impact of an industry average alkaline battery, based on input from four battery manufacturers. The whole life cycle of the battery is established as a first goal of the study to provide a context for these end‐of‐life impacts. The second section, described in the remainder of the report (Chapter 5 & 6) focuses just on the end of life treatment for alkaline batteries. The goal of this second part of the study is to compare different disposal and recycling scenarios for alkaline batteries to weigh the environmental burdens and benefits of each specific situation for battery disposition at end of life. More detail around scope for the second section will be established at the beginning of Chapter 5. The intended audience of the study is NEMA, local and state government agencies including waste management entities, as well as the general environmental community through journal publication. The geographic region of interest for product sales and use is the United States and portions of Canada. System boundaries are defined in accordance with the ecoinvent life cycle inventory database3 unless otherwise specified to include all life‐cycle steps from material extraction to end‐of‐life (this database presents some limitations because of its EU focus, given the US geographic scope of this study). For the first section of the study the cut‐off for EOL materials to recycling is applied.
Transport Transport Transport … Raw Material 1 Raw Material 2 Raw Material X Transport Retailers Manufac‐ turing
Use End‐of‐Life Production
Materials Production Not includedstudy in Distribution
FIGURE 5. PHASES WITHIN THE PRODUCT LIFE CYCLE.
The terminology for the phases within the product life cycle is defined in Figure 5. At the highest level, the life cycle is broken into three phases: production, use, and end‐of‐life. The production phase is further broken into manufacturing (i.e., battery manufacturing), packaging, and distribution. The environmental impacts of distribution centers and retailers are ignored in this analysis, but the transportation to them is included. Finally, the manufacturing phase is further broken down into the production of raw materials (i.e. acquisition and refining of raw materials) and the actual manufacturing of the battery. Transportation of raw materials from suppliers to the manufacturing facility is included in materials production. For the purposes of this study the use phase contributes no environmental impact because the investigation focuses only on primary alkaline batteries. For primary batteries the beneficial work they do for the consumer in use derives from the chemical potential of the materials contained within the device. As stated previously, the end‐of‐life (EoL) phase provides the primary area of interest for this particular investigation in the second section of the study. Regardless of the scenario under investigation at EoL the batteries go through some intermediate transportation and consolidation phase and are then taken to a particular disposal facility. The EoL boundaries and scenarios will be fully outlined after the full life cycle results are presented. The functional unit used in this analysis is 1 kg weighted average alkaline batteries, equivalent to 30 batteries; this quantity was chosen as most relevant to a consumer. This weighted average is the sales weighted average of the batteries sizes sold in the US (as described in detail in Chapter 4). For the full life cycle assessment described in chapter 4, impacts are shown of 1 kg of these weighted average batteries including packaging and for the end‐of‐ life analysis the functional unit is the treatment of 1 kg of batteries.
The data for the study will be gathered through the battery manufacturers participating, the potential battery recyclers, literature, modelling efforts and interviews with collection system providers. Therefore the data are of varying quality and are required to be transparent to the research team for interpretation. Specifics on the temporal, geographic and representativeness of these data are provided within the chapters below. The critical review process for this study will be done by three external reviewers sequentially from three different institutions in academia and state government. Those reviews can be found in the appendices. The interpretation of this study involves identifying the
significant issues found from the inventory and analysis steps, evaluating the completeness of the work and providing a description of the gaps and recommendations. As mentioned above, the specific impact assessment methods used in this study were: cumulative energy demand, global warming potential and the three damage categories of Ecoindicator 99. Value choices are made by selecting just these five metrics to present in the results. Future work should look at other characterization approaches.
LKALINE BATTERY LIFE CYCLE ASSESSMENT
This section describes the full life cycle analysis for the primary alkaline battery to provide the relevant scale of end‐of‐life compared to the rest of the life cycle. This full alkaline battery LCA is not directly applicable to the second study goal but provides important context. There are other reasons to focus on end‐of‐life besides its impact on the whole battery life cycle. The set of parameters outlined in the upcoming tables are used for all analyses in this section.
For the alkaline battery life cycle assessment, each phase of the life cycle is identified. Following this, materials and energy are quantified and environmental impacts are calculated for each phase. This section describes the methodology in detail by identifying the scope of the analysis for the alkaline battery and then describing the sources of data – including necessary assumptions. The data used in this analysis was gathered through a survey of the firms participating in the study and then averaged by a statistician within NEMA. The survey used is provided in Appendix A.
As described in Figure 5, the life cycle consists of production, use, and a standard end‐of‐life treatment. The production phase for the alkaline battery consists of producing the raw materials, transporting the raw materials to the manufacturing facility, manufacturing the battery, transporting the battery to the packaging facility, packaging the battery, transporting the packaged battery to distribution facilities throughout the United States, and finally transporting the packaged batteries to retail facilities. It is assumed that no environmental impact is associated with the use phase of the alkaline battery because it is single use and any emissions to air, land or soil in this use of a battery in a product would be attributed to that product. The end of life treatment for this first section is just taken as standard landfill and incineration without any materials recovery. More detail around end of life is investigated in the second section of the study. For the allocation of recycling materials at end of life (for example in the manufacturing process), the cut off approach is used, so no credit or burden is assigned to the portion of material recycled in that process at end of life. While a scenario analysis on this could be performed, the overall impact of the recycled scrap after manufacturing provides a small amount of the overall impact. The same is true for the recycled packaging at end of life.
A single alkaline battery is actually represented as the weighted average of each size of battery (AA, AAA, C, D and 9V) based on percentage sales in 2007 as shown in Table 2. This weighted average is used
to determine the weight of the battery, the bill of materials, the amount of packaging and the weighted distance traveled in each transport step. Finally, the baseline end‐of‐life scenario assumed for the alkaline battery consists of landfilling and incineration (87% landfill and 13% incineration) . TABLE 2. ALKALINE BATTERY SALES BROKEN DOWN BY SIZE FOR 2007 2007 Sales Weight of battery (g) AA 60% 23 AAA 24% 11 C 4% 71 D 8% 147 9V 4% 45
DATA SOURCES AND ASSUMPTIONS
This section describes the data gathered and processed for implementation. It should be noted that the data shown here do not represent a company‐specific battery bill of materials or manufacturing process but instead are an aggregation of data from four separate OEMs. First, the bill of materials is established for a single weighted‐average alkaline battery. The mass of a single weighted average alkaline battery (WAAB) is 33 grams. The six top components by mass and each component’s mass within the battery calculated from the weighted mass percentages of battery constituents and these 33 grams are shown in Table 3. The chemistry of each of the sizes of alkaline battery (AA, AAA, C, D and 9V) are assumed to be the same (although the weight percentage of materials for the 9V is slightly different). Line 2 contains 35wt% aqueous potassium hydroxide for the electrolyte.
Table 3 also shows the supplier locations for each material (by country) and the one‐way distance traveled from that supplier to the manufacturing facility by truck and boat (backhaul distances were ignored). A few supplier locations for each material were provided and the distances reported in the table are the average of each country listed. For overseas shipments, specific ports‐of‐call were assumed when they were not specified. There was no information provided about manufacturing scrap, so the partial bill of materials shown below is the actual amount of material in the weighted battery. However, a few exceptions to this include 1) it was difficult to determine the amount of water present in the electrolyte as received (described below) and 2) the remaining mass was divided by the “other materials” (brass, plastics [incl. PVC], paper, and galvanized steel) .
TABLE 3. TOP TEN COMPONENTS BY MASS WITHIN THE BILL OF MATERIALS FOR A SINGLE WAAB.
No. Material Mass
(g) Supplier Locations Nominal Distance from Supplier Truck (km) Boat (km) 1 Electrolytic Manganese Dioxide 13 Japan, South Africa, US 1600 11000 2 Potassium hydroxide (35wt% aqueous KOH) 3.7 US 1000 ‐ 3 Graphite 1.2 Brazil, Canada, Switzerland 1130 5900 4 Nickel‐Plated Steel 6.0 Japan, Netherlands, US 1600 7700 5 Zinc 5.8 Canada, Japan, US 1450 9000 6 Brass 1.0 ‐ ‐ Not specified 7 Galvanized steel 0.52 8 Nylon 0.51 9 Paper 0.51 10 PVC 0.51 Total weighted battery 33 Also included as an input in the analysis but not shown in the bill of materials are the excess materials needed in production that ends up as scrap (as reported consisting primarily of steel). Information on the additional material, water and energy inputs and waste outputs (including the scrap materials) from the manufacturing facility were also provided, through the survey, in units of input or output per million weighted average batteries (based on production within the facility). The values for these inputs and outputs from manufacturing are shown in Table 4, allocated by unit to a single WAAB. The inputs to the facility were electricity, natural gas and light fuel oil as well as water. Outputs from the facility were also provided, including the waste for recycling steel, waste to landfill, water treatment, and air emissions in the form of volatile organic compounds.
It is common industry practice that the water used in production is to dilute the as‐received 50% potassium hydroxide to the concentration used in the final electrolyte (35%). This water is included in the bill of materials. The incoming water used in production that is not accounted for in the bill of materials in Table 3 is accounted for as an input to the facility and output of wastewater leaving the site.
TABLE 4. INPUTS AND OUTPUTS FROM THE BATTERY MANUFACTURING FACILITY ALLOCATED TO A SINGLE WAAB. Inputs Amount per Battery Units Water 32 g Electricity 0.02 kWh Natural Gas, burned in industrial furnace 25 kJ Fuel Oil, burned in industrial furnace 9.3 kJ Outputs Amount per Battery Units VOC 0.02 g Waste (for recycling) 1.1 g Waste (for disposal) 0.52 g Waste Water 32 g *The number for VOC’s in this inventory is observed to be high, but this was as reported from the company survey described above. This turns out to be a minimal contribution to the results.
The next phase of the life cycle is packaging of the battery. Packaging occurs at a facility 460 km from manufacturing (weighted distance by sales). The materials used to package the batteries in “blister packs” of 2 or 4 (depending on the size), as shown in Table 5, are polyvinyl chloride, paperboard and corrugated cardboard for shipping. Approximately 6% of the final mass of the packaged battery is attributed to these packaging materials.
TABLE 5. MATERIALS USED IN THE PACKAGING OF A SINGLE WAAB.
No. Material Amount per
Battery Units 1 Polyvinyl Chloride 0.4 g 2 Corrugated board 1 g 3 Paper board 0.8 g
The inputs and outputs from the overall operation of the packaging facility are provided in Table 6 allocated to a single WAAB. TABLE 6. INPUTS AND OUTPUTS FROM THE PACKAGING FACILITY ALLOCATED FOR A SINGLE WAAB. Inputs Amount per Battery Units Water 1.7 g Electricity 5.5 Wh Natural Gas, burned in industrial furnace 6.1 kJ Outputs Amount per Battery Units Waste Water 1.7 g
After the batteries are packaged they are shipped to the distribution centers. The nominal distance for transport from packaging to the distribution centers (based again on the weighted average of sales) is 630 km. In addition, 1100 km is used as a weighted average distance from the distribution centers to the retailer. Table 7 provides a summary of each segment of transport from manufacturing to distribution. The mass of the battery up until the packaging facility is 33 g; after packaging, the mass of the battery and packaging is approximately 35 g in total. TABLE 7. TRANSPORTATION FOR ALKALINE BATTERY FROM MANUFACTURING TO PACKAGING AND FINALLY DISTRIBUTION.
From To Nominal
Manufacturing Packaging 460 Truck
Packaging Distribution Center 630 Truck
Distribution Center Retailer 1100 Truck
As mentioned previously, there is no additional environmental impact from the use phase of the alkaline battery since no energy is added beyond the production of the cell detailed above. The final phase in the life cycle is the end‐of‐life treatment of the battery. This analysis assumed that 13% of alkaline batteries are incinerated at end‐of‐life as that is the percent of total generation of MSW that is combusted in the US . The remaining percentage of alkaline batteries is landfilled; both scenarios assumed to involve 100 km of MSW vehicle transport. A generic landfilling and incineration scenario is used for this baseline case, just to understand the order of magnitude comparison between production and EoL. Table 8 summarizes the baseline end‐of‐life scenario described above for the battery. TABLE 8. END OF LIFE DESCRIPTION FOR THE ALKALINE BATTERY. Transport Distance (km) Disposal Truck 100 Waste Scenario Percent Landfill 87% Incineration 13% Table 9 summarizes the end‐of‐life scenario for the battery packaging, which involves recycling of 30% of the cardboard packaging and the disposal of the remaining packaging through landfill and incineration. Because of the assumption of cut off allocation at end of life no burden or benefit is associated with the quantity of recycled packaging material; a reduced end of life burden is assumed.
TABLE 9. END OF LIFE DESCRIPTION FOR THE ALKALINE BATTERY PACKAGING. Transport Distance (km) Recycle Truck 400 Disposal Truck 100 Material Recycled Percent Cardboard 30% Waste Scenario Percent Landfill 87% Incineration 13%
ECOINVENT DATA GAPS
The source of data for implementation of this portion of the analysis is the ecoinvent 2.2 database. A few major assumptions were necessary for implementation and they are mentioned here, especially in the case when an inventory is not available for a particular item in the bill of materials. The first major limitation of ecoinvent is its European focus, while the geographic scope of this study was the US. One gap in the ecoinvent database was an inventory for manganese dioxide, the major component in the alkaline bill of materials (~40 wt% or 13 g in a WAAB). Previous studies, for example the Defra‐ commissioned report mentioned earlier, have substituted the manganese inventory for manganese dioxide, altering the mass stoichiometrically . Substituting an inventory for titanium dioxide for manganese dioxide is another possible approach, as both materials can be produced using similar processes. Manganese dioxide is produced through a series of steps, including roasting of manganese ore, dissolution of the roasted manganese in acid, filtration of impurities, and the electrowinning of the final product. The inventory for manganese(III) oxide developed for lithium ion batteries (Mn2O3) was used as a proxy. The steel inventory used includes 40% steel from an electric arc furnace and the remainder converted pig iron from a blast furnace. The zinc inventory includes 30% from combined zinc production. The full outline of the inventories used is provided in Appendix B.
A 16‐32‐tonne truck is assumed to perform all land transport, while all boat transport is performed by transoceanic freight ship, both from the ecoinvent database.
Where primary data was not available secondary sources were used, including published reports, specifications, and the ecoinvent LCI database. The quality of the data has been assessed on the following criteria:
• Source‐‐primary or secondary
• Representativeness—how closely the data collected represents the supply chain of the system, including geographic and operational considerations
Stage Data Source Temporal Representativeness
Alkaline Battery Materials Primary data regarding types of material and quantities. Secondary data (ecoinvent) for upstream extraction and processing Primary data from 2009, based on ecoinvent processes (global mix used when available), Zinc and steel inventories are EU focused Limits on data based on global focus of ecoinvent inventories Battery Primary Packaging Primary data regarding types of material, and quantities. Secondary data (ecoinvent) for upstream extraction and processing Primary data from 2009, other based on ecoinvent processes Limits on data based on European focus of ecoinvent data Battery Manufacturing Facility Primary data on energy consumption quantities and waste produced. Secondary data (ecoinvent) for inventory, electricity grid mix for US used. Primary data from 2009, other based on ecoinvent processes Limits on data based on EU focus of some inventories, however electricity mix US‐ based Packaging Facility Primary data regarding energy consumption quantities and waste produced. Secondary data (ecoinvent) for inventory Primary data from 2009, other based on ecoinvent processes Limits on data based on EU focus of some inventories, however electricity mix US‐ based Transportation Primary data regarding transportation distances Secondary data (ecoinvent) for inventory Primary data from 2009, other based on ecoinvent processes Limits on data based on European focus of ecoinvent data 50% load factor truck likely overestimate of burden as trucks are likely more full. Use No environmental burden associated with use as the chemical energy stored in the battery through use of manganese dioxide and zinc Disposal (for this section) Secondary data (ecoinvent) for disposal scenarios. Data does not well represent the geographic scope
The following section describes the results from the full life cycle analysis for 1 kg WAAB including their packaging where the end of life fate is 87% landfill, 13% incineration w/o steel recovery as this is investigated in more detail in the second section of this work. Results describe the impact assessments from the life cycle inventory using Cumulative Energy Demand (CED) in MJ, Global Warming Potential (GWP) in grams of CO2 equivalent, and Eco‐Indicator (EI) midpoints for Human health (in DALY),
Ecosystem quality (in PDF*m2yr) and Resources (MJ Surplus). The methods differ in focus, as CED emphasizes total energy consumption, GWP stresses global warming contributing gases, and EI may highlight perceived human health risks (Human health indicator) or ecosystem toxicity (Ecosystem quality). Furthermore, the final EI damage category, Resources, comments on perceived resource scarcity of a particular material or the upstream impacts associated with that material. In some cases only the results of one impact assessment method are presented. The beginning of this analysis explores the “hot spots” for impact within the life cycle of an alkaline battery, resolving where the biggest impacts are. The full Eco‐Indicator 99 breakdown is provided at the end of this section. Figure 6 shows the relative contribution of each phase on the full life cycle impact and plots the values in for CED and Figure 7 shows the relative contribution of each phase for GWP. FIGURE 6. LIFE CYCLE IMPACT USING CED FOR 1 KG WAAB INCLUDING PACKAGING(~30 BATTERIES) FIGURE 7. LIFE CYCLE IMPACT USING GWP FOR 1 KG WAAB INCLUDING PACKAGING (~30 BATTERIES)
Table 10 shows the life cycle impact of 1 kg WAAB using all three LCIA methodologies, CED, GWP and the three damage categories of EI 99. Table 10, Figure 6, and Figure 7 show that the production phase dominates the life cycle impact. TABLE 10. LIFE CYCLE IMPACT OF 1 KG WAAB, INCLUDING PACKAGING USING THREE LCIA ASSESSMENT METHODS (THE DAMAGE CATEGORIES OF EI 99 ARE SHOWN, HUMAN HEALTH, ECOSYSTEM QUALITY AND RESOURCES). Life Cycle Phase CED (MJ/1 kg WAAB) GWP (kg CO 2 eq./1 kg WAAB) Human Health (DALY/1 kg WAAB) Ecosystem Quality (PDF*m2yr/1 kg WAAB) Resources (MJ surplus/1 kg WAAB) Production 66 3.8 1.1x10‐5 1.5 4.7 End‐of‐Life 2.5 0.6 9.7x10‐7 0.62 0.18 TOTAL 68 4.3 1.2x10‐5 2.1 4.9 % EoL contribution 4% 13% 8% 29% 4% Table 1 is also repeated below in Table 11 but with the impact of one WAAB and 1 kg WAABs included to provide some perspective for the impacts shown in these results (both include packaging). TABLE 11. LIFE CYCLE IMPACT ASSESSMENT VALUES AS SHOWN IN TABLE 1 WITH THE ADDITION OF ONE WEIGHTED‐AVERAGE ALKALINE AND 1 KG WEIGHTED AVERAGE ALKALINE BATTERIES.
Product or Process CED
(MJ) GWP (kg CO2 eq) Human Health (DALY) Ecosystem Quality (PDF*m2yr) Resources (MJ surplus) Production of 25 g PET beverage bottle (20 fl oz/590 ml) 2 0.07 6.8 x 10 ‐8 0.005 0.15 1 weighted‐average battery (33 g) 2 0.14 4 x 10‐7 0.07 0.16 Production of 14 g aluminum beverage can (12 fl oz/350 ml) 3 0.2 2.6 x 10 ‐7 0.007 0.17 1 kg weighted‐average batteries 68 4.3 1.2 x 10‐5 2.1 4.9 100 km fuel consumption in a European passenger car 305 18 2 x 10 ‐5 1 21 Coffee pot: 5 years 5400 220 2 x 10‐4 30 190
GWP, Human Health and Ecosystem Quality have a higher relative contribution from the end‐of‐life scenario than CED and Resources resulting from the generic landfilling and incineration processes. These will be examined in much more detail in the subsequent analysis.
Because the production phase dominates the life cycle, drilling further down into the production phase reveals the drivers of impact within that phase. Table 12 shows the values of this breakdown for 1 kg of WAAB. Figure 8 shows the absolute values using CED plotted side by side.
TABLE 12. BREAKDOWN OF PRODUCTION IMPACTS FOR 1 KG OF WA ALKALINE BATTERIES USING FIVE INDICATORS (DOES NOT INCLUDE END‐OF‐LIFE, WHICH WAS PRESENTED IN TABLE 11)
Life Cycle Phase
CED (MJ/1 kg WAAB) GWP (kg CO 2 eq./1 kg WAAB) Human Health (DALY/1 kg WAAB) Ecosystem Quality (PDF*m2yr/1 kg WAAB) Resources (MJ surplus/1 kg WAAB) Materials Production 43 2.5 9.3x10 ‐5 1.4 3.4 Manufacturing 10 0.6 5.6x10‐7 0.022 0.55 Transport 6.9 0.42 5.1x10‐7 0.037 0.51 Packaging materials 5.6 0.22 2.2x10‐7 0.04 0.24 TOTAL 66 3.7 9.4x10‐5 1.5 4.7 FIGURE 8. BREAKDOWN OF PRODUCTION IMPACTS FOR 1 KG WAAB USING CED.
There are a few differences between the indicators used, which will be discussed later in this section. However, all three indicators show that the impacts of materials production dominate production
The results so far have combined the transport of raw materials from the suppliers to the manufacturing facility with the production of the raw materials themselves in the category ‘materials production’. The manufacturing of the raw materials dominates this impact, as transportation is a very small percentage of the burden.
Figure 9 shows the ranking of materials within materials production for CED. This ranking changes depending on which indicator is used as shown in Table 13 (while the uncertainty in the results makes it difficult to differentiate between the materials in the last several places in the table, the first few materials are much different in impact even with the attendant uncertainty). In CED and GWP, and Resources manganese dioxide has the largest impact; in EI, the zinc has the largest impact. This reflects the perceived toxicity of zinc by this metric. Figure 9 demonstrates how quickly the impact falls off after the first three materials. The bulk of the burden is thus focused on three materials: manganese dioxide, zinc ingot, and steel. FIGURE 9. MATERIALS PRODUCTION IMPACTS FOR 1 KG WAAB USING CED (THESE VALUES INCLUDE THE TRANSPORT OF RAW MATERIALS TO THE MANUFACTURING FACILITY).
TABLE 13. RANKING OF ENVIRONMENTAL BURDEN FOR MATERIALS FOR FIVE IMPACT METHODS CED (MJ/1kg WAAB) GWP (kg CO2 eq/1kg WAAB) Human Health (DALY/1kg WAAB) Ecosystem Quality (PDF*m2yr/1kg WAAB) Resources (MJ surplus/1kg WAAB) Mass in 1kg WAAB (g)
MnO2: 17 MnO2: 1.1 Brass: 3.4x10‐6 Zinc: 0.9 MnO2: 1.1 MnO2: 390
Zinc: 9.8 Zinc: 0.52 Zinc: 2.7x10‐6 Brass: 0.25 Steel: 0.76 Steel*: 190
Steel: 7.5 Steel: 0.46 MnO2: 1.2x10‐6 Steel: 0.12 Brass: 0.6 Zinc**: 170
Nylon: 1.8 Nylon: 0.11 Steel: 1.0x10‐7 MnO2: 0.071 Zinc: 0.5 KOH***: 110
KOH: 1.8 KOH: 0.094 Nickel: 8.3x10‐7 Nickel: 0.036 Nickel: 0.17 Graphite: 36
Brass: 1.3 Brass: 0.04 KOH: 1.0x10‐7 Paper: 0.02 Nylon: 0.13 Brass: 31
PVC: 1 Nickel: 0.047 Nylon: 7.6x10‐8 KOH: 0.006 KOH: 0.1 Paper:15
Paper: 1 PVC: 0.033 Paper: 2.5x10‐8 Nylon: 0.0018 PVC: 0.062 Nylon: 15
Nickel: 0.8 Paper: 0.014 PVC: 2.0x10‐8 Graphite: 0.001 Paper: 0.014 PVC: 15
Graphite: 0.2 Graphite: 0.011 Graphite: 1.7x10‐8 PVC: 0.0008 Graphite: 0.01 Nickel: 3.6
*Includes steel in can and galvanized steel;**Just zinc in electrode and galvanized steel, not brass;***KOH including water
This analysis shows that for CED, GWP, and resources, the greatest environmental impact of alkaline batteries comes from the materials production of manganese dioxide. For all three of these metrics, approximately 1/3 of the total environmental impact from production comes from a single material. This general trend mirrors the highest materials by mass within the battery. However, for the case of the human health and ecosystem quality indicator, zinc has the highest environmental burden, reflecting the relative toxicity of zinc in human health and the ecosystem as described by this method. Brass also comes to the forefront for similar reasons in human health and ecosystem quality indicator. Neither of these components is highest by mass, but because of their relative perceived toxicity, they come to the top of the list for ecosystem quality or toxicity. In the ecoinvent inventory process zinc emissions to air in the primary production process (from mining and processing) dominates the environmental impact associated with both human health and ecosystem quality, accounting for ~90% of the burden of zinc. Finally, it is useful to examine the burden of the manufacturing facility in a bit more detail. Table 14 and Figure 10 demonstrate the relative impact of electricity, natural gas, diesel, water and waste in the battery manufacturing facility.
TABLE 14. RELATIVE CONTRIBUTIONS OF IMPACTS OF THE MANUFACTURING FACILITY FOR 1 KG WAAB. Manufacturing Facility CED (MJ/1 kg WAAB) GWP (kg CO 2 eq./1 kg WAAB) Human Health (DALY/1 kg WAAB) Ecosystem Quality (PDF*m2yr/1 kg WAAB) Resources (MJ surplus/1 kg WAAB) Electricity 8.9 0.53 5.3x10‐7 2.1x10‐2 0.45 Natural Gas 0.9 0.051 1.6x10‐8 4.1x10‐4 0.071 Diesel 0.36 0.024 1.1x10‐8 1.4x10‐3 0.028 Water 0.018 7.7x10‐4 1.0x10‐9 6.6x10‐5 7x10‐4 Waste 0.0086 4.8x10‐4 1.1x10‐9 6.3x10‐5 5.5x10‐4 TOTAL 10 0.6 5.5x10‐7 2.3x10‐2 0.5 FIGURE 10. BREAKDOWN OF THE IMPACTS OF MANUFACTURING FOR 1 KG WAAB USING CED.
From Figure 10 it is clear that the electricity use within has the greatest effect on the environmental burden of the manufacturing facility. Another element of note in Figure 10 is the negative value (or credit) associated with the waste and recycling burden (indicated by the black bar below the y‐axis). This credit is associated with the steel recycling within the manufacturing facility.
As described above there is no environmental burden from the use phase of an alkaline battery. The impact of the end‐of‐life scenario consisting of 87% landfilling and 13% incineration was a small portion of the overall impact as described in Table 10. The EOL scenario will be further investigated in the remainder of this report.
As a final set of summarizing analyses for this full life cycle assessment, two more details around the Eco‐Indicator 99 metric are presented. Table 15 outlines the specific results of each characterization for a 1 kg WAAB weighted alkaline battery.
TABLE 15. A) PRODUCTION VERSUS END‐OF‐LIFE IMPACTS AND B) WITHIN RAW MATERIALS, PRODUCTION, TRANSPORTATION AND PACKAGING FOR EACH CHARACTERIZATION WITHIN ECO‐INVENT FOR 1 KG WAAB.
Impact Units Production End‐of‐Life
DALY/1 kg WAAB
Respiratory organics 3.4E‐09 8.4E‐10
Respiratory inorganics 3.9E‐06 1.7E‐07
Climate change 7.9E‐07 1.1E‐07
Radiation 2.2E‐08 3.0E‐10
Ozone layer 2.6E‐10 2.5E‐11
Ecotoxicity PDF*m2yr/ 1 kg WAAB 1.3 0.61 Acidification/Eutrophication 0.1 0.006 Land use 0.1 0.003 Minerals MJ surplus/ 1 kg WAAB 1.1 0.001 Fossil fuel use 3.5 0.180
Impact Units Materials Prod. Manufacturing Transportation Packaging
DALY/1 kg WAAB
5.7E‐06 1.3E‐07 3.5E‐08 5.5E‐08
Respiratory organics 2.1E‐09 5.2E‐10 5.7E‐10 2.4E‐10
Respiratory inorganics 3.1E‐06 2.9E‐07 3.8E‐07 1.3E‐07
Climate change 5.2E‐07 1.3E‐07 8.8E‐08 5.7E‐08
Radiation 1.7E‐08 3.3E‐09 7.8E‐10 1.2E‐09
Ozone layer 1.6E‐10 2.5E‐11 6.9E‐11 1.2E‐11
Ecotoxicity PDF*m2yr/ 1 kg WAAB 1.3 8.6E‐03 1.3E‐02 0.005 Acidification/Eutrophication 0.1 0.0094 1.8E‐02 0.0042 Land use 0.1 0.0047 6.6E‐03 0.033 Minerals MJ surplus/ 1 kg WAAB 1.1 0.0033 6.1E‐03 0.0028 Fossil fuel use 2.3 0.55 5.0E‐01 0.25
To summarize the full life cycle implications of alkaline batteries, the production of raw materials dominates the life cycle with the transport of those raw materials to manufacturing having a minimal environmental impact. A few materials dominate this materials production impact, with manganese dioxide, zinc, and steel having the highest relative impacts. To return to the data quality assessment for this phase of the analysis, despite the ubiquity of European data based on ecoinvent, the relative impact of materials and elements of manufacturing are assumed to be reliable. Furthermore, within manufacturing the electricity burden dominated which was modeled after a US grid, therefore also more relevant geographically. Therefore the data gaps, in particular around geography, are not expected to have an effect on the dominant elements of the alkaline battery life cycle.